Bioresource Technology 171 (2014) 175–181

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Methane and nitrous oxide emissions following anaerobic digestion of sludge in Japanese sewage treatment facilities Kazuyuki Oshita a,c,⇑, Takuya Okumura b, Masaki Takaoka a, Takashi Fujimori a, Lise Appels c, Raf Dewil c a

Graduate School of Global Environmental Studies, Kyoto University, C-cluster, Kyotodaigaku-katsura, Nishikyo-ku, Kyoto 615-8540, Japan Graduate School of Engineering, Kyoto University, C-cluster, Kyotodaigaku-katsura, Nishikyo-ku, Kyoto 615-8540, Japan c Department of Chemical Engineering, KU Leuven, Jan De Nayerlaan 5, 2860 Sint-Katelijne-Waver, Belgium b

h i g h l i g h t s  CH4 and N2O emissions after anaerobic sludge digestion were investigated. 3

 CH4 mission factor was 509 ± 72 mg/m -influent and especially high in winter. 3

 N2O emission factor was 7.1 ± 2.6 mg/m -influent and lower than CH4 emission.  The highest CH4 emissions were during dewatering, followed by continued digestion.  CH4 and N2O emissions after anaerobic digestion are considered to be significant.

a r t i c l e

i n f o

Article history: Received 15 June 2014 Received in revised form 14 August 2014 Accepted 17 August 2014 Available online 23 August 2014 Keywords: Sewage sludge Anaerobic digestion Methane Nitrous oxide Dewatering

a b s t r a c t Methane (CH4) and nitrous oxide (N2O) are potent greenhouse gases with global warming potentials (expressed in terms of CO2-equivalents) of 28 and 265, respectively. When emitted to the atmosphere, they significantly contribute to climate change. It was previously suggested that in wastewater treatment facilities that apply anaerobic sludge digestion, CH4 continues to be emitted from digested sludge after leaving the anaerobic digester. This paper studies the CH4 and N2O emissions from anaerobically digested sludge in the subsequent sludge treatment steps. Two full-scale treatment plants were monitored over a 1-year period. Average emissions of CH4 and N2O were 509 ± 72 mg/m3-influent (wastewater) and 7.1 ± 2.6 mg/m3-influent, respectively. These values accounted for 22.4 ± 3.8% of the indirect reduction in CO2-emissions when electricity was generated using biogas. They are considered to be significant. Ó 2014 Elsevier Ltd. All rights reserved.

1. Introduction The anaerobic digestion of sewage sludge produces biogas consisting mainly of methane (CH4), which can be used as an energy source. However, the continuation of CH4-emissions to the atmosphere from digested sludge during the subsequent sludge treatment steps may significantly contribute to global warming. Indeed, the 100-year global warming potential of CH4 is 28, meaning that each emitted molecule of CH4 contributes 28 times more to global warming than an emitted carbon dioxide (CO2) molecule during this period (IPCC, 2013). CH4 produced by sludge is, moreover, combustible, hence giving rise to fire and explosion hazards (Astbury, 2008). Whereas CH4 emissions from composting and wetlands have ⇑ Corresponding author at: Graduate School of Global Environmental Studies, Kyoto University, C-cluster, Kyotodaigaku-katsura, Nishikyo-ku, Kyoto 615-8540, Japan. Tel.: +81 75 383 3336; fax: +81 75 383 3338. E-mail address: [email protected] (K. Oshita). http://dx.doi.org/10.1016/j.biortech.2014.08.081 0960-8524/Ó 2014 Elsevier Ltd. All rights reserved.

been reported before in some papers (e.g., Maulini-Duran et al., 2013; Uggetti et al., 2012), little information is available on the emission of CH4 during the dewatering process after anaerobic digestion (IPCC, 2002). Additionally, nitrous oxide (N2O) has an even higher 100-year global warming potential of 265 (IPCC, 2013), and is released as a by-product of the nitrification and/or denitrification processes. Here again, only very few studies are available on N2O emissions following the anaerobic digestion of sludge (Czepiel et al., 1995; Scherson et al., 2014; Toyoda et al., 2011; Tallec et al., 2008). In Japan, the current emission factor for CH4 from sewage sludge dewatering amounts 106 mg/m3-influent (wastewater) (Ministry of the Environment, Japan, 2006). This factor, which is frequently used in estimations for greenhouse gas inventories, was determined from an investigation of treatment plants 20 years ago (Sato et al., 1992; Takeishi et al., 1993) and is based on two assumptions: (i) an anaerobic sludge digestion process is installed before the sludge dewatering, and (ii) all of the dissolved CH4 in

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digested sludge is emitted during sludge dewatering. Because of these assumptions, all CH4 generated in the sludge between the anaerobic digestion and the sludge dewatering process was included in the current emission factor. Also, the obtained value of 106 mg/m3-influent is not applicable for sewage treatment plants in which anaerobic sludge digestion is not employed. A N2O emission factor for sludge dewatering is not yet available (Ministry of the Environment, Japan, 2006). Given the contributions of CH4 and N2O emissions to global warming, the production of these gases after anaerobic sludge digestion may considerably reduce the environmental benefit of using biogas as an energy source to indirectly reduce CO2 emissions. A reliable quantification of the CH4 and N2O emissions between the anaerobic digestion and sludge dewatering processes is necessary to be able to calculate and take into account these effects. Several recent studies reporting life-cycle assessments on sludge treatment (including anaerobic digestion), did not take into account this aspect and only examined the fugitive biogas emissions by incomplete combustion and failure of combustion (Brown et al., 2010; Cao and Pawłowski, 2013; Patterson et al., 2011; Venkatesh and Elmi, 2013). In Europe, Daelman et al. (2012) concluded that CH4 generated by the wastewater and sludge treatment processes has a high global warming potential in CO2 equivalents, and that its impact on global warming reduces the indirect reduction in CO2 emissions derived from biogas with sewage treatment as a fuel source. The present study quantified the emissions of CH4 and N2O from the anaerobic digestion and sludge dewatering processes and used these data to update the emission factors in Japan. More specific, gaseous CH4 and N2O emissions in the unit operations related to the sludge dewatering process were measured (these process steps where these gases are most likely to be emitted). Also the water removed from the sludge during the dewatering process was analyzed for dissolved CH4 and N2O. The CH4 and N2O emission factors and their influence on the indirect reductions in CO2 emissions derived from using biogas as a fuel source at two sewage treatment facilities over a 1-year period were also quantified. 2. Methods 2.1. Sewage treatment facilities Flow diagrams for the sewage treatment facilities and locations of sample collection/measurement points are provided in Fig. 1. Two sewage facilities were selected for this study: one facility employing anaerobic sludge digestion (H) and one without (N). The total amount of wastewater annually treated in the two facilities was very similar: 38,005,000 m3/year and 38,745,489 m3/year for facilities H and N, respectively (Japan Sewage Works Association, 2010). Information on the wastewater composition in both facilities throughout the measurement campaign (between April 2011 and March 2014) is provided in the Supplementary information (Table S1). Although, total solids (TS), volatile total solids (VTS), suspended solids (SS) and biochemical oxygen demand (BOD) of the influent in facility H were higher than in facility N, the total nitrogen concentrations and relative composition of the influent in both facilities were comparable. In facility H, the produced sludge is subjected to a mesophilic anaerobic digestion [hydraulic retention time (HRT): 36 days; average temperature: 39 °C; organic matter decomposition ratio: 64.6%] after thickening. Most of the biogas generated during anaerobic digestion is used as auxiliary fuel for sludge incineration. The digested sludge is dewatered using a belt-filter press, producing a sludge cake containing 84% water. The dewatered sludge is then combusted in a fluidized bed incinerator. The off-gas from the dewatering process is deodorized using activated carbon in the exhaust ventilation system. In facility N, the sludge is directly dewatered

(without prior anaerobic digestion) using a screw-type dewatering (achieving a sludge cake with a water content of 79%), also followed by fluidized bed incinerator. The off-gas from the dewatering process is biologically deodorized in the exhaust ventilation system. 2.2. Sampling point and dates Sampling was done at those points in the sewage treatment facility that represent potential sources of CH4 and N2O (Fig. 1). Facility H uses an open-type storage tank for the digested sludge, and CH4 emissions from the surface of the digested sludge were measured using the floating closed-chamber method (Czepiel et al., 1993). CH4 and N2O concentrations were also measured in the off-gas and in the water separated from the sludge during the dewatering process. In facility N (not applying anaerobic digestion), CH4 and N2O were measured in the off-gas and in the water separated from the sludge during the dewatering process. Odorous off-gas from the storage tank of the sludge cake is used as combustion air in the incineration chamber, hence this gas stream was not included in the measurements as all components are further oxidized during the combustion. Also, samples of digested sludge in facility H and thickened sludge in facility N were collected and their CH4 and N2O potentials were compared. On-site sampling and measurements were carried out during two winter and two summer periods: November 2011–March 2012 (winter), August– September 2012 (summer), July–September 2013 (summer), and December 2013–January 2014 (winter). 2.3. Measurements of CH4 in the digested sludge storage tank Flux emissions of CH4 from the storage tank of digested sludge in facility H were measured using the floating closed chamber method (Figs. S2 and S3 in the Supplementary information), consisting of a cylindrical case made of polyvinyl chloride (PVC) and a float around the case composed of polyethylene foam. Two ports were located on the upper side of the case; one of these was used for gas sampling, and the other was attached to a 1-L gas bag inside the case to maintain constant air pressure. Measurements of CH4 were carried out during daytime for a total of 8 days in winter and 6 days in summer. The weather was clear on all measurement dates. The floating closed chamber was used at 4–6 points on the sludge surface in the storage tank during 1 day. The temperature inside the chamber was recorded, and gas emitted from the sludge was continuously analyzed for CH4 concentration using a CH4 gas detector (XP-3160; New Cosmos Electric Co., Ltd., Osaka, Japan); the gas flow rate and sampling time were 200 mL/min and 5 min, respectively. Measurements of CH4 were used to calculate rates of CH4 production (dC/dt), and CH4 fluxes (lmol/m2/s) from the surface of the digested sludge, according to Czepiel et al. (1993). The method of calculation is included in Supplementary information (Section S.3). N2O emissions were also measured, but all measurements were below the detection limit and, therefore, will not be discussed further. 2.4. Measurement of CH4 and N2O in water separated from sludge during the dewatering process The water separated from the sludge during the dewatering process in both sewage treatment facilities was analyzed for CH4 and N2O. Sampling was done over the course of 6 days in winter and 3 days in summer. The analytical procedures outlined in Hatamoto et al. (2010) were followed for quantifying CH4 and N2O dissolved in water separated from sludge during the dewatering process. Specifically, water samples were collected in 122 mL vials by overflowing the

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Fig. 1. Flow diagrams for sewage treatment facilities (H, N) and sampling points (black boxes with white text) for measurements of methane (CH4) and nitrous oxide (N2O).

volume of the container to prevent exposure of the sample to air. Vials were capped with butyl rubber septa and sealed with aluminum caps using a hand-crimper. A portion of the water sample (30 mL) was subsequently withdrawn from the vial using a syringe and replaced with 30 mL of nitrogen gas (N2). A disinfectant (10 mL of 20% chlorhexidine gluconate solution) was added to the vial to inhibit biological oxidation and generation of CH4 and N2O. Sample vials were placed in a water-bath shaker (T-2S; Thomas Kagaku Co., Ltd., Tokyo, Japan) maintained at 40 °C and shaken at 100 rpm for 1 h to achieve a gas–liquid equilibrium for CH4 and N2O. Gas samples were extracted from the vials using a 50-mL syringe, and CH4 and N2O were quantified using a gas chromatograph (Micro-GC CP2002 and CP4900; Agilent Technologies, Inc., Santa Clara, CA, USA). The concentration of CH4 and N2O in the water samples was calculated according to Hatamoto et al. (2010), Weiss and Price (1980), and Wiesenburg and Guinasso (1979). The fill calculation method is provided in the Supplementary information (Section S.4). 2.5. Sample collection and measurements of CH4 and N2O in the offgases from the dewatering process Samples of the off-gases were collected over 8 days during winter and 6 days during summer in facility H, and over 11 days during winter and 6 days during summer in facility N. Samples were collected after the deodorizing process (facility H: activated carbon; facility N: biological deodorization). 1 L samples were taken of the off-gases of the dewatering process using gas bags. Samples were collected four times over the course of 1 day in each sewage treatment facility. Off-gas flow rates were measured in-line following the International Standard (ISO 5801, 2007). Concentrations of CH4 and N2O were quantified using a Micro-GC, and total CH4 and N2O emissions were subsequently calculated by multiplying the off-gas flow rate by their respective concentrations.

procedures (APHA, 1992). Sludge (80 mL) was collected in 122mL vials, which were capped with butyl rubber septa and sealed with aluminum caps using a hand-crimper. Vials were placed in a water-bath shaker kept at 40 °C and shaken at 50–60 rpm. The volume of gas generated in the headspace of the vial was measured using a syringe, and the concentrations of CH4 and N2O in the headspace were quantified using a Micro-GC in a period between 6 h and 30 days. Of the sludge samples collected, 2–3 samples of each type of sludge were used to determine the gas production and CH4 and N2O concentrations over time. 3. Results and discussion 3.1. Measurements of CH4 flux Measurements of CH4 flux and air temperature for both seasons are presented in Fig. 2. The CH4 flux (arithmetic mean ± 1 standard error) in facility H was 30.2 ± 3.84 lmol/m2/s (n = 34) and 46.2 ± 8.86 lmol/m2/s (n = 31) during the winter and summer seasons, respectively. No significant difference in CH4 flux was observed between both seasons (p > 0.05; Student t-test). The CH4 emissions per unit volume of digested sludge was 1.53 ± 0.19 mol/m3 digested

9.7

30.9

2.6. CH4 and N2O generation potential of sludge The production potentials of digested sludge (facility H) and thickened sludge (facility N) for CH4 and N2O were investigated. The total solid (TS) and volatile total solid (VTS) content of sludge samples were analyzed in triplicate according to standard

Fig. 2. Arithmetic means of air temperature (top) and methane (CH4) flux from the digested sludge storage tank during winter (n = 34) and summer (n = 31) in facility H. Bars represent 1 standard error, and * indicates p > 0.05.

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sludge (n = 65), calculated based on the open surface of the storage tank (78.5 m2) and the rate of digested sludge entering the tank (7 m3/h). The theoretical saturation point for dissolved CH4 (MCH4) was calculated as follows:

M CH4 ¼

e  S39  C CH4  1000 d39

ð1Þ

where e is the water pressure placed on the digested sludge withdrawn from the bottom of the digestion tank (2.2 atm), calculated from the height of the digestion tank (12 m); S39 is the Bunsen solubility coefficient of CH4 (0.0256, v/v; 39 °C, 1 atm) calculated by Wiesenburg and Guinasso (1979); CCH4 is the concentration of CH4 in biogas (60 vol.%); and d39 is the volume per mol of CH4 at 39 °C and 1 atm (25.6 L/mol). The calculation resulted in a value of 1.32 mol/m3-H2O for MCH4. This is slightly lower than the measured value of 1.53 ± 0.19 mol/ m3 for the digested sludge. The difference between the calculated and measured values could be explained by the supersaturation of the digested sludge by CH4 (biogas present in small gas bubbles in the digested sludge) and/or a continued anaerobic degradation of organic matter in the storage tank. Digested sludge was supplied to the storage tank from a pipe located 500 mm above the surface of stored digested sludge, hence it is expected that a major fraction of the CH4 dissolved in the digested sludge was already emitted to the atmosphere during introduction. This was confirmed by the slight difference in CH4 flux between summer and winter. A CH4 emission factor of 119 ± 15 mg/m3-influent was calculated based on the annual sewage treatment volumes (38,005,000 m3/year) and digested sludge production (185,220 t/year; Japan Sewage Works Association, 2010). This emission factor is similar to the current CH4 emission factor for sludge dewatering in Japan, which assumes that all dissolved CH4 in digested sludge is emitted during the dewatering process. 3.2. Measurements of CH4 and N2O in water separated from sludge during the dewatering process Fig. 3 shows the arithmetic mean concentrations and 1 standard error for the dissolved CH4 and N2O in the water separated from the sludge during the dewatering process, for both seasons and facilities. The concentrations of dissolved CH4 were lower in facility H than in facility N (p < 0.05; Student t-test). This can be explained due to differences in the amount of water used in the washing process applied in the dewatering device in each facility: the amount of water used to wash the sludge cake in the belt-filter press in facility H was higher than that for the screw-press in facility

Fig. 3. Arithmetic mean of methane (CH4) and nitrous oxide (N2O) concentrations in water separated from sludge during the dewatering process in facilities H and N. Bars represent 1 standard error. Numbers on the x-axis denote the number of samples (n).

N (US EPA, 2000). Obviously, this washing water dilutes the water released from the sludge and has an influence on the CH4 concentration measured. Concentrations of dissolved CH4 were higher in the winter than in the summer in facility N (p < 0.05; Student t-test), which may be due to increased water solubility at lower water temperatures. The concentrations of dissolved N2O were an order of magnitude lower than the concentrations of dissolved CH4, despite the fact that N2O is more soluble in water. Although very little dissolved N2O was detected in facility H, the concentrations of dissolved N2O in facility N were higher, especially in summer. This difference may be due to differences in the volume of water released during the dewatering process. However, this observation can also be related to the generated and dissolved N2O at the installed sewage and sludge treatment processes. Firstly, since the treated volume, nitrogen concentration and main composition of the influent at facility H and N were comparable (Table S1), and the type of installed biological process (a conventional activated sludge process in Facility H and an anaerobic/anoxic/oxic (A2O) process in facility N) was previously shown not to have a significant influence on total amount of N2O generated from the sewage treatment, differences will only occur due to the applied process conditions. E.g. the aeration rate, SRT and the organic loading rate and nitrification stage are dominant process parameters, whereas the denitrification stage is less important (Kampschreur et al., 2009). Moreover, N2O can be generated during sludge treatment (including dewatering) as a byproduct of occurring denitrification if nitrite is present in the sludge (Kampschreur et al., 2009). Both previously cites mechanisms may affect the dissolved N2O concentrations in facility N. Emission factors calculated for CH4 from water separated from the sludge were 11.4 ± 6.44 mg/m3-influent and 0.47 ± 0.21 mg/m3-influent for winter and summer, respectively, in facility H, and 19.4 ± 1.81 mg/m3-influent and 2.86 ± 0.12 mg/m3-influent for winter and summer, respectively, in facility N. The calculations were based on the annual volume of water separated from the sludge in facilities H (159,919 m3/year) and N (102,904 m3/year) (Japan Sewage Works Association, 2010). These factors were much lower than the CH4 emission factor for the digested sludge itself (119 ± 15 mg/m3-influent, as previously reported in the manuscript). Emission factors calculated for N2O from water separated from sludge were 0.0117 ± 0.005 mg/m3-influent and 0.0151 ± 0.005 mg/m3-influent for winter and summer, respectively, in facility H and 1.14 ± 0.086 mg/m3-influent and 0.350 ± 0.150 mg/m3-influent for winter and summer, respectively in facility N. 3.3. Measurements of CH4 and N2O in deodorized off gas generated from the dewatering process The arithmetic mean and 1 standard error of the CH4 and N2O concentrations measured in the deodorized off-gas generated from the dewatering process in both sewage facilities are presented in Fig. 4. The off-gas flow rates were 60.5 m3/min and 61.3 m3/min in facilities H and N, respectively. The CH4 concentration in the off-gas during winter was seven times higher than the concentration in summer in facility H (p < 0.05; Student t-test). This difference was most likely due to the anaerobic decomposition of organic matter in sludge prior to anaerobic digestion during summer. Following anaerobic digestion, very little CH4 will be generated between the digested sludge storage tank and the dewatering process. To confirm this hypothesis, quantification of CH4 emissions from sludge treatment processes preceding anaerobic digestion (i.e., the thickening process) are necessary in future studies. The concentration of CH4 in facility N, which does not employ anaerobic digestion, was lower than in facility H; the CH4 concentration in facility N was particularly lower in winter than in summer (p < 0.05; Student

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need to be revised, and a distinction should be made between sewage treatment with and without anaerobic digestion. 3.4. CH4 and N2O generation potential of sludge

Fig. 4. Arithmetic mean of methane (CH4) and nitrous oxide (N2O) concentrations in deodorized off-gas generated from the dewatering process in facilities H and N. Bars represent 1 standard error. The numbers on the x-axis indicate the number of samples (n).

t-test), possibly due to the increased air temperature during summer. Emission factors calculated for CH4 from the dewatering process were 667 ± 89.9 mg/m3-influent and 99.7 ± 17.2 mg/m3-influent in winter and summer, respectively, in facility H, and 127 ± 24.7 mg/m3-influent and 219 ± 23.2 mg/m3-influent in winter and summer, respectively, in facility N based on the off-gas flow rates. These factors are comparable to (facility H) or considerably higher (facility N) than the current CH4 emission factor for dewatering (106 mg/m3-influent). The concentration of N2O in facility N was higher than that in facility H. This difference is not likely due to differences in the volume of water used in the washing process during dewatering, but is attributable to N2O generated by the nitrification at the sewage treatment process and/or the denitrification at the sludge treatment process. Emission factors calculated for N2O were below the detection limit and 14.1 ± 5.23 mg/m3-influent in winter and summer, respectively, in facility H, and 33.3 ± 7.41 mg/m3-influent and 23.1 ± 4.25 mg/m3-influent in winter and summer, respectively, in facility N. The emission factors calculated for CH4 and N2O indicate that the dewatering process significantly contributes to the total CH4 and N2O emissions, which is especially the case for CH4 in winter, when more gas is generated between the digested sludge storage tank and dewatering processes. These results indicate that the current CH4 and N2O emission factors for sludge treatment in Japan

Fig. 5. Volume of methane (CH4) generated from sludge (at 40 °C) in each facility and season. Each plot is the arithmetic mean (n = 17–27). Bars represent 1 standard error.

Since CH4 can be generated between the sludge digestion and dewatering processes, the potential of digested sludge (facility H) and thickened sludge (facility N) to generate CH4 and N2O in both seasons was investigated. The results are presented in Fig. 5. In facility H with a digester, the amount of CH4 generated from the digested sludge in both seasons was 100 mL/g VTS, with a solids retention time (SRT) of 30 days, which was obviously lower than the volume of CH4 generated from raw sludge in a typical digestion tank (240–300 mL/g VTS; Appels et al., 2008). Digested sludge still has the potential to generate CH4, as the volume and generation rate of CH4 from digested sludge at the beginning of experiment were higher in winter than in summer. Since the SRT of the storage tank and dewatering process is 17 h, the calculated CH4 amounts from the results of Fig. 5 were 10.3 mL/g VTS in winter and 6.9 mL/g VTS in summer. These results are consistent with the observed trend for CH4 emission factors from the dewatering process, which were 667 ± 89.9 mg/m3-influent in winter and 99.7 ± 17.2 mg/m3-influent in summer. In facility N, which does not have a digester, the volume and rate of CH4 production from dewatered sludge were low and exhibited a 4-day lag-phase in gas generation. This lag-phase is commonly encountered and represents the time required for physiological and regulatory processes to adapt to new environments (Rolfe et al., 2012). The low volume and rate of CH4 production in facility N is, therefore explained by the low amount of anaerobic microorganisms present in thickened sludge. However, this result contradicts the fact that the CH4 emission factors in facility N were 127 ± 24.7 mg/m3-influent in winter and 219 ± 23.2 mg/m3-influent in summer, and considerably more than the facility’s CH4 emission factor in summer (99.7 ± 17.2 mg/m3-influent). These higher values should be explained by the anaerobic digestion of sludge accumulated in the dead volume between the sludge pipe and dewatering machine. Most measurements of N2O generated from sludge were below the detection limit, but N2O was detected in the deodorized off-gas at the sludge dewatering in Facility H in summer. This N2O originates from newly generated N2O after digestion by the denitrification and is emitted by ventilation of the dewatering device. 3.5. Quantitative influence of CH4 and N2O emissions following anaerobic digestion on the indirect reduction in CO2 emissions by biogas generation The arithmetic mean of CH4 and N2O emission factors in summer and winter are provided in Table 1. To convert the CH4 and N2O emission factors to CO2-equivalent emission factors, ratios of 28 and 265 were used (IPCC, 2013). The largest source of CO2-equivalent CH4 emission in facility H was deodorized off-gas generated from the dewatering process, followed by continued anaerobic digestion. The largest source of CO2-equivalent N2O emission in facility N was deodorized off-gas generated from the dewatering process, which was greater than the contribution to total CH4 emissions. Emissions of N2O were mostly affected by the nitrification at sewage treatment and/or the denitrification during sludge treatment including dewatering process, which requires further investigation. CO2-equivalent emission factors for CH4 and N2O following anaerobic digestion were compared with indirect reductions in CO2 emissions from using biogas as a fuel in facility H. The following scenarios were evaluated in facility H: (A) emissions of CH4 and N2O following digestion were excluded; (B) current emission

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Table 1 Emission factors of methane (CH4) and nitrous oxide (N2O) and their carbon dioxide (CO2) equivalent in this study. Gas

Facility

Storage tank of digested sludge

Deodorized ventilation off-gases from dewatering building

Dissolved gas in separated water from dewatering

Total

Total: CO2 equivalent

CH4

H N H N

119 ± 15 – N.D. –

384 ± 53.6 173 ± 23.9 7.0 ± 2.6 28.2 ± 5.8

5.9 ± 3.3 11.1 ± 1.0 0.01 ± 0.005 0.7 ± 0.1

509 ± 72 184 ± 25 7.1 ± 2.6 29.0 ± 5.9

14.3 ± 2.0 5.2 ± 0.7 1.9 ± 0.7 7.7 ± 1.6

N2O

Unit: mg/m3-influent (only CO2 equivalent value: g-CO2 eq/m3-influent). –: not installed; N.D.: not detected.

factors for CH4 and N2O from the dewatering process were adopted (106 mg/m3-influent, 0 mg/m3-influent); and (C) CH4 and N2O emission factors for facility H (Table 1) were adopted. The indirect reduction in CO2 emissions (Kd; g CO2 eq/m3-influent) was calculated as follows:

K d ¼ f d  pd  E=W  1000

ð2Þ

where Fd is the annual volume of biogas produced (2,649,800 m3N/year) (Japan Sewage Works Association, 2010), Pd is the electricity generated by a gas engine fueled with only biogas per unit volume of biogas (1.88 kWh/m3N used in all scenarios) (actual value of facility H), E is the amount of CO2 produced per unit of electricity (0.555 kg CO2/kWh) (Horio et al., 2009), and W is the annual volume of influent (38,005,000 m3-influent/year) (Japan Sewage Works Association, 2010). Given these parameters, a value of 72.2 g CO2 eq/m3-influent was calculated for Kd. The CO2-equivalent CH4 and N2O emission factors following anaerobic digestion in scenario N (N = A, B or C) (KeN; g CO2 eq/ m3-influent) were calculated as follows:

K eN ¼ ðMN  GWPm þ NN  GWP n Þ=1000

ð3Þ

where MN (N = A, B or C) represents the mass of CH4 produced per unit volume of influent, with a value of 106 mg/m3-influent used in scenario B (MB) and 509 ± 72 mg/m3-influent used in scenario C (MC); GWPm is the global warming potential of CH4 (28); NC is the mass of N2O produced per unit volume of influent in scenario C (7.1 ± 2.6 mg/m3-influent); and GWPn is the global warming potential of N2O (265). The calculated results, KeA, KeB and KeC were 0.0, 3.0 and 16.1 ± 2.7 (g CO2 eq/m3-influent), respectively. If all biogas generated in the digestion tank is directly emitted to the atmosphere, the CO2-equivalent CH4 in biogas emission factors calculated by Fd, W and GWPm was 837 (g CO2 eq/m3-influent) and much higher than KeB and KeC. (KeB/Kd) and (KeC/Kd) were 0.041 and 0.224 ± 0.038 respectively. The contribution of CO2-equivalent CH4 and N2O emissions following anaerobic digestion to the indirect reduction in CO2 emissions as a result of biogas generation was 4.1% and 22.4 ± 3.8% in scenarios B and C, respectively. The contribution in scenario C, which was based on the results of this study, was higher than expected compared to scenario B. These values were also higher than the established fugitive losses of biogas due to incomplete combustion and failure of combustion, which are estimated to be 0.3–1.0% (IPCC, 2006; US EPA, 2009). Daelman et al. (2012) calculated that the CO2-equivalent of CH4 emissions contributed 46.5% to the indirect reduction in CO2, but it is unclear why this value is significantly larger than the results of the present study. However, it is clear from the results of this study that CH4 emissions following anaerobic digestion have a considerable negative effect on the indirect reduction in CO2 by using biogas as a fuel. Further research is needed to determine whether the results of this study reflect the patterns of CH4 and N2O emissions in other sewage treatment facilities, including the effects of the nitrification

at sewage treatment and/or the denitrification sludge treatment on N2O emissions. Actions to limit the CH4 production following anaerobic digestion that can be implemented and are considered effective include (i) the thermal and SRT control of CH4 fermentation so that digestion is more complete in the digestion tank, and (ii) decreasing CH4 fermentation activity by reducing the transit time of digested sludge to the dewatering process or by cooling the transport pipe from the sludge digestion tank to the dewatering site with effluent water. As for the CH4 generated in deodorized off-gas from the dewatering process, the use of this gas as combustion air in the sludge incinerator is limited because of the large gas volume and the very low CH4 concentration. Therefore, using an energy-saving method is necessary to decompose the rarefied CH4 in deodorized off-gas. CH4 in deodorized off-gas might be decomposed if the off-gas is sent into the aeration tank as aeration air (Daelman et al., 2012). 4. Conclusions This study investigated CH4 and N2O emissions following the anaerobic digestion of sewage sludge. CH4 emissions were higher in winter in facility H due to continued anaerobic digestion of sludge. The emission factors for CH4 and N2O were 509 ± 72 mg/ m3-influent and 7.1 ± 2.6 mg/m3-influent, respectively, in the sewage treatment facility with a digester (H). The contribution of CO2-equivalent CH4 and N2O emissions following anaerobic digestion to the indirect reduction in CO2 emissions resulting from using biogas as a fuel was 22.4 ± 3.8% and is considered to be significant. Acknowledgements We thank the staff members of sewage treatment facility H and N who supported us in the sampling process. Part of this study was supported financially by a grant for strategic international research network program for vitalizing brain circulation from the Japan Society for the Promotion of Science (JSPS).

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